Environmental Toxicology

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1 Environmental Toxicology and Chemistry, Vol. 32, No. 4, pp , 2013 # 2013 SETAC Printed in the USA DOI: /etc.2110 Environmental Toxicology DIETARY EXPOSURE OF MINK (MUSTELA VISON) TO FISH FROM THE UPPER HUDSON RIVER, NEW YORK, USA: EFFECTS ON REPRODUCTION AND OFFSPRING GROWTH AND MORTALITY STEVEN J. BURSIAN,*yz JOHN KERN, RICHARD E. REMINGTON, JANE E. LINK,y and SCOTT D. FITZGERALDk# ydepartment of Animal Science, Michigan State University, East Lansing, Michigan, USA zcenter for Integrative Toxicology, Michigan State University, East Lansing, Michigan, USA KERN Statistical Services, Sauk Rapids, Minnesota, USA kdepartment of Pathobiology and Diagnostic Investigation, Michigan State University, East Lansing, Michigan, USA #Diagnostic Center for Population and Animal Health, Michigan State University, Lansing, Michigan, USA (Submitted 31 July 2012; Returned for Revision 3 September 2012; Accepted 27 September 2012) Abstract The effects of feeding farm-raised mink (Mustela vison) diets containing polychlorinated biphenyl (PCB) contaminated fish from the upper Hudson River (New York, USA) on adult reproductive performance and kit growth and mortality were evaluated. Diets contained 2.5 to 20% Hudson River fish, providing 0.72 to 6.1 mg P PCBs/g feed ( pg toxic equivalents [TEQ WHO 2005 ]/g feed). The percentage of stillborn kits per litter was significantly increased by dietary concentrations of 4.5 mg P PCBs/g feed (28 pg TEQ WHO 2005 /g feed) and greater. All offspring exposed to dietary concentrations of 4.5 and 6.1 mg P PCBs/g feed (28 and 38 pg TEQ WHO 2005 /g feed) died by 10 weeks of age, and all offspring exposed to 1.5 and 2.8 mg P PCBs/g feed (10 and 18 pg TEQ WHO 2005 /g feed) died by 31 weeks of age, leaving juveniles in the control and 0.72 mg P PCBs/g feed (0.41- and 4.8 pg TEQ WHO 2005 /g feed) groups only. The dietary concentration predicted to result in 20% kit mortality (LC20) at six weeks of age was 0.34 mg P PCBs/g feed (2.6 pg TEQ WHO 2005 /g feed). The corresponding maternal hepatic concentration was 0.80 mg P PCBs/g liver, wet weight (13 pg TEQ WHO 2005 / g liver, wet wt). Mink residing in the upper Hudson River would be expected to consume species of fish that contain an average of 4.0 mg P PCBs/g tissue. Thus, a daily diet composed of less than 10% Hudson River fish could provide a dietary concentration of P PCBs that resulted in 20% kit mortality in the present study. Environ. Toxicol. Chem. 2013;32: # 2013 SETAC Keywords Hudson River Mink Polychlorinated biphenyl Reproduction Mortality INTRODUCTION The Hudson River is contaminated with polychlorinated biphenyls (PCBs) from Fort Edward, New York, USA, to New York City. Capacitor manufacturing facilities at Fort Edward and Hudson Falls, New York, USA, are considered to be the major sources of PCBs in the upper Hudson River, with discharges beginning in Between 1966 and 1974, the Fort Edward and Hudson Falls facilities purchased 35,000 metric tons of PCBs or 15% of domestic sales in the United States. It is estimated that approximately 600 metric tons of PCBs were released into the Hudson River between the 1940s and 1977 [1,2]. Concentrations of PCBs measured in the livers of mink (Mustela vison) collected in the vicinity of the Hudson River are above thresholds associated with effects and have not decreased over time. Mink collected over 25 years ago had hepatic concentrations of PCBs [3] that were equivalent to concentrations associated with reproductive impairment in ranch mink experimentally exposed to PCBs [4,5]. More recently, mink collected in the vicinity of the Hudson River 16 years after the initial study [3] had no apparent decrease in hepatic PCB concentrations [6], which were above the criteria for impairment of mink health and reproduction [7,8]. Mink are among the most sensitive species to PCBs [9 18]. Because mink are fish-eating mammals that satisfy criteria, All Supplemental Data may be found in the online version of this article. * To whom correspondence may be addressed (bursian@msu.edu). Published online 4 January 2013 in Wiley Online Library (wileyonlinelibrary.com). including chemical sensitivity, for a sentinel wildlife species [19], regulatory agencies such as the U.S. Environmental Protection Agency (U.S. EPA), U.S. Fish and Wildlife Service, Environment Canada, and the Swedish Environmental Protection Agency recognize the mink as such, making it one of the most commonly selected receptors in ecological risk assessments for sites involving aquatic habitats with elevated concentrations of PCBs and related compounds [20]. Thus, the objective of the present study was to evaluate the health effects of feeding farm-raised mink diets containing PCB-contaminated fish from the Hudson River. In the present report, the effects on adult reproductive performance (percentage of females whelping, gestation length, percentage of stillbirths, and live kits whelped) and offspring growth and mortality through 31 weeks of age are discussed in terms of dietary and hepatic concentrations of P PCBs and World Health Organization (WHO) 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) toxic equivalents (TEQs WHO 2005 ). A companion paper presents the effects on organ mass and pathology, including the mandible and maxilla, of adult mink and their offspring [21; this issue]. METHODS Fish preparation Common carp (Cyprinus carpio) and Atlantic herring (Clupea harengus) were used as fish components of the experimental diets. Carp were collected from the upper Hudson River from Northumberland Pool (114 kg), from the vicinity of Lock 2 (879 kg), and from the first 610 m of Moses Kill (528 kg; Fig. 1) 780

2 Effects of Hudson River fish on mink reproduction Environ. Toxicol. Chem. 32, Fig. 1. Map indicating the three sites on the upper Hudson River (New York, USA) from which carp (Cyprinus carpio) was collected. by the New York Department of Environmental Conservation and transported frozen to the Michigan State University Experimental Fur Farm. Fish were ground, blended, sampled for contaminant and nutrient analyses, and frozen for subsequent incorporation into mink feed. Whole Atlantic herring, purchased from Finicky Pet Food, was chosen as the control fish because they, like carp, contain the enzyme thiaminase [22]. The herring was ground, blended, and sampled as described above. Dietary treatments Treatment diets were based on the Michigan State University Experimental Fur Farm ranch diet formulated to meet the nutrient requirements of mink [22]. Diets contained 25% whole ground chicken, 24% water, 20% fish, 15% wheat middlings (Akey), 4% spray-dried eggs (Van Eldren), 3.5% spray-dried liver (Van Eldren), 3.5% soybean oil (North American Nutrition), 3% spray-dried blood (California Spray Dry), 1% phosphoric acid (food grade, 75%; Astaris), 0.48% vitamin premix (Akey), 0.48% mineral premix (Akey), and 0.04% biotin (Archer Daniels Midland). The proportion of fish incorporated into the feed (20%) is in the range of fish consumed by mink in the wild [12]. The control diet contained 20% ocean herring, and the remaining five treatment diets contained a mixture of the herring and Hudson River fish in the following proportions: 2.5% Hudson River fish/17.5% herring; 5% Hudson River fish/ 15.0% herring; 10% Hudson River fish/10% herring; 15% Hudson River fish/5% herring; 20% Hudson River fish. The decision to incorporate no more than 20% Hudson River fish into the diet was based on the high mean P PCB concentration in the collected fish (36 mg P PCBs/g wet wt) and the desire to not cause complete reproductive failure and/or adult mortality. Targeted dietary concentrations of P PCBs were 0, 0.90, 1.8, 3.6, 5.4, and 7.2 mg/g feed based on 36 mg P PCBs/g wet weight in the Hudson River fish. Analysis Dietary P PCBs. Dietary and tissue samples were analyzed for a number of chemicals. Samples of each treatment diet were shipped to Alpha Woods Hole Laboratory for analysis of P PCBs, PCB homolog groups, and potentially toxic and bioaccumulative metals and minerals. For the PCB analysis [23], samples were homogenized in diatomaceous earth containing 1:1 dichloromethane:acetone and exchanged to hexane. The extracts were cleaned with sulfuric acid. The extract was then analyzed for PCB congeners and homolog groups by lowresolution gas chromatography mass spectrometry (GC MS) in selective ion monitoring (SIM) mode. Blanks, duplicates, and spikes were analyzed at a frequency of one per analytical batch of up to 20 samples. In addition, to evaluate accurate and consistent extraction and analysis, a standard reference material (SRM), SRM 1946 (National Institute of Standards and Technology, USA, Lake Superior Fish Tissue), was extracted and analyzed with each analytical batch. This SRM has certified values for 30 PCB congeners. Results for the congener analyses were to be within 20% of the 95% confidence interval (CI) for certified concentrations. Overall, greater than 93% of the reference material results were within the control limits for the certified concentrations. Dietary metals. Individual metals (Supplemental Data, Table S1) were analyzed by U.S. EPA method 6020A [24] using inductively coupled plasma-mass spectrometry. Mercury was analyzed by U.S. EPA method 7471A [25] using cold vapor atomic absorption spectroscopy. Dietary and tissue PCB congeners. Diet samples were also sent to Axys Analytical Services for analysis of PCB congeners,

3 782 Environ. Toxicol. Chem. 32, 2013 S.J. Bursian et al. polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs), chlorinated pesticides, and polybrominated diphenyl ethers (PBDEs). In addition, liver samples from adults, six-week-old kits, and 31-week-old juveniles were submitted for analysis of PCB congeners. For chlorinated pesticides (Supplemental Data, Table S1) and PCB congeners, the sample was homogenized with anhydrous powdered sodium sulfate and extracted with dichloromethane. The extracts were cleaned with a Biobeads SX-3 (Bio-Rad) gel permeation column, followed by a Florisil (US Silica) column cleanup [26]. The extract was then split, and one split was analyzed for pesticides [27] and most PCB congeners [26] using low-resolution GC MS. The other split was analyzed for coplanar PCBs [26] with high-resolution GC MS. The split for coplanar PCBs was further cleaned using an acid/base-layered silica column, a 4.5% carbon AX-21 (Anderson Development)/celite 545 (Sigma-Aldrich) mixture, and finally an alumina column. Dietary and tissue PCDD, PCDF, and PBDE congeners.for PCDD/PCDF and PBDE analyses, the procedure was the same through the Biobeads SX-3 gel permeation column. After that, a fluid-management system was used with the following order of columns: jumbo acid silica, layered acid/base silica, alumina, and a carbon column. Individual PCDD/PCDF congeners [28] and individual PBDE congeners [29] were analyzed using highresolution GC MS. The percentage of lipids was determined gravimetrically for all samples by the weight of the total residue contained in an aliquot of the sample extract. Dietary and hepatic concentrations of analytes are presented on a wetweight basis unless otherwise noted. Data validation. Data validation was based on the qualityassurance/quality-control criteria documented in the Analytical Quality Assurance Plan: Hudson River Natural Resource Damage Assessment (Version 2.0, September 1, 2005), U.S. EPA National Functional Guidelines for Organic Data Review (1999), U.S. EPA National Functional Guidelines for Inorganic Data Review (1994), and the individual laboratory standard operating procedures cited above. Animals The Michigan State University Institutional Animal Care and Use Committee approved the use of animals in the present study. Seventy-five first-year (virgin), natural dark, female mink and 30 first-year, natural dark, male mink from the Michigan State University Experimental Fur Farm herd were assigned to the six treatment groups on December 26, Littermates were not placed in the same treatment group, to minimize genetic predisposition to PCB toxicity. The control and targeted 5.4- and 7.2 mg P PCBs/g feed groups had 15 females and five males each and the targeted 0.90-, 1.8-, and 3.6 mg P PCBs/g feed groups had 10 females and five males each. The number of mink placed on trial balanced a reasonable expectation of detecting biologically meaningful events subject to the limitations of available time and resources to conduct the study. The control and the two highest treatment groups were assigned 15 females rather than 10 to increase sample sizes at dietary concentrations where severe effects were anticipated. Housing Female mink were housed individually in suspended wire cages (76 cm L 61 cm W 46 cm H) on the outside aisles of an open-sided mink shed. Five animals per treatment group were assigned to a bank of five cages. Assignment of treatments to banks of cages was done randomly. A wooden nest box (38 cm L 28 cm W 27 cm H) bedded with aspen shavings and excelsior (wood wool) was attached to the outside of each cage. Males were assigned to banks of five cages (61 cm L 30 cm W 38 cm H) on an inside aisle of the same shed. Feed and water were available ad libitum. The Standard Guidelines for the Operation of Mink Farms in the United States [30] were followed for housing and maintenance of animals. Exposure period Mink were started on their respective treatment diets on January 3, 2007, after a one-week acclimation period. Fresh feed was provided daily, and water was available ad libitum. One hour prior to feeding the treatment diets each day, animals were given 10 g of control feed to which thiamine had been added (0.25 mg thiamine hydrochloride/animal) to prevent Chastek s paralysis that could result from the incorporation of thiaminase-containing fish into the feed [22]. Feed consumption was determined for the same 2-d period each week, and body mass was recorded every two weeks. On February 14, 2007, in response to a decrease in feed consumption in the groups fed the diets that contained predominantly herring, the decision was made to incorporate a flavor enhancer (Proliant B3301 Spray-dried beef flavor) and additional vitamin E into all treatment diets at rates of 66 and 4.4 g/kg feed, respectively. Animals were provided the diets containing the flavor enhancer beginning February 15, 2007, through the end of the trial. Measurement of feed consumption and determination of body mass were discontinued at the initiation of breeding, to reduce stress to the females. Measurement of feed consumption was not resumed after whelping because the females and their litters consumed the same feed until weaning. After weaning, kits were group-housed until they were approximately 10 weeks old, making accurate determination of individual feed consumption difficult. Because adult feed consumption cannot be measured from initiation of breeding through weaning of kits, the value for average daily feed consumption determined for the first seven weeks of the trial was applied for the total duration of the trial. Females were mated to males within their respective treatment groups between March 1 and March 21, Each female was given an opportunity to mate every fourth day until a presumed successful mating occurred as determined by posture during mating and postcoital appearance of the female s vaginal area. A mated female was given an opportunity to breed with a different male the day following a successful mating and on the eighth and ninth days after a successful mating (a common commercial mink breeding practice). The whelping period began on April 22 and ended on May 9, Nest boxes were checked on a daily basis for the presence of mink kits. The gender of the kits was determined at birth, and the live and stillborn young were counted. Body mass of kits was recorded at birth and at three and six weeks of age. Body mass of adult females was recorded at the time their litters were weighed. All surviving adult mink were killed with CO 2 and necropsied between June 6 and June 19, Adult males were necropsied on June 6, 2007, adult females that did not whelp or lost their kits before weaning were necropsied on June 7, 2007, and adult females that whelped and successfully weaned their kits were necropsied on June 12 and June 19, 2007, with the exception of one female in the 7.2 mg P PCBs/g feed group that was still nursing a kit. The necropsy dates were chosen to allow assessment of placental scarring in females that did not whelp and in females that lost their kits before the uterine horns

4 Effects of Hudson River fish on mink reproduction Environ. Toxicol. Chem. 32, regressed and to allow those females that had kits to continue nursing until kits were six weeks of age. Altogether, 14 control females, nine 0.90 mg P PCBs/g feed females, nine 1.8 mg P PCBs/g feed females, two 3.6 mg P PCBs/g feed females, and three 5.4 mg P PCBs/g feed females were killed and necropsied during the last phase. The single surviving female in the 7.2 mg P PCBs/g feed group was necropsied on June 28, On June 12 and June 19, 2007, a sample of weanling kits (approximately six weeks old) was killed and necropsied as previously described. This sample included 15 control kits (seven females, eight males), nine 0.90 mg P PCBs/g feed kits (four females, five males), and mg P PCBs/g feed kits (seven females, six males). There were not sufficient numbers of kits in the 3.6-, 5.4-, and 7.2 mg P PCBs/g feed groups to justify necropsy of weanlings. Sampling was done to insure that as many litters as possible were represented. The remaining kits were maintained on their respective treatment diets. On July 17, 2007, littermates were separated, with animals being assigned to individual cages (61 cm L 30 cm W 38 cm H) in the two interior rows of the open-sided shed. Body mass was determined on a monthly basis until termination of the trial during the first week in December. On December 4 and 5, 2007, seven-month-old animals in the two remaining treatment groups (16 females and 14 males in the control group and 11 females and 12 males in the 0.90 mg P PCBs/g feed treatment group) were killed and necropsied as previously described. Statistical analyses Toxic equivalents were calculated by summing the products of individual PCDD, PCDF, and PCB congener concentrations and their respective toxic equivalency factors (TEF) [31]. Congeners that had concentrations less than the detection limit were assigned a concentration equal to one-half the detection limit. The choice of assigned value (0, one-half detection limit, or detection limit) had no substantive influence on P PCB congener or TEQ concentrations based on a quantitative assessment. Measurement end points of interest were classified into one of three data types and statistically analyzed according to data type. Possible data types were as follows: continuous measurements, such as SPCBs in livers; counts, such as kits per litter; or binary outcomes, such as whether or not an individual kit survived to three weeks. Statistical analyses of measurement end points were conducted using a generalized linear model framework [32] where the most appropriate class of linear models was selected based on classification of data type and correlation structure (e.g., repeated measures, kits clustered within litters). A summary of endpoints classified by data type and analysis method is provided in Table 1. Continuous endpoints were analyzed by analysis of variance (ANOVA) [33] when the endpoint was measured at the experimental unit level and the experimental units within a treatment group were expected to be independent (e.g., adult mink). For example, gestation lengths for pregnant females in the same treatment group were expected to be independent. Continuous endpoints having repeated measures on an individual animal or kits clustered within litters were analyzed with linear generalized estimating equation (GEE) regression. The GEE models are a contemporary extension of generalized linear models for clustered data that adjust for within-cluster correlation [34]. Examples of clustered data in the present study were repeated measures of adult female body mass and kit liver SPCB concentrations within litters. The adult female body mass model for the prebreeding period was adjusted for baseline body mass and days on treatment. Likewise, the adult female feed consumption model was adjusted for days on treatment. Count endpoints measured on adults, including number of kits whelped and kits whelped alive, were analyzed using negative binomial regression models. Treatment effects on the rate of kits whelped alive were estimated as differences (%) in rate of kits whelped alive per litter from control. Refitting the model with SPCBs in feed as a continuous variable, the relative rate of live kits per litter for a given increase in micrograms P PCBs/g feed (D) was estimated. The minimum dietary concentrations necessary to induce 20 and 50% kit mortality (LC20 and LC50) were estimated based on the maximum likelihood estimates provided by the betabinomial regression. Dose response relationships were estimated for SPCBs as well as TEQs. The LC20 and LC50 values expressed as SPCBs and TEQs were estimated (with 95% CIs) for stillbirth, mortality at three weeks, and mortality at six weeks of age. Control-adjusted LC20 and LC50 values were also estimated based on the same maximum likelihood estimates of the beta-binomial regression and summarized in Supplemental Data, Table S2. An LC50 expressed as dietary SPCB and TEQ concentrations for mortality by 31 weeks of age was estimated as the concentration corresponding to 1 minus the joint probability of survival from birth to the six-week necropsy period and from the six-week necropsy period to the end of the juvenile growth trial (31 weeks of age). The product of these survival Table 1. Summary of study end points, data types, and statistical methods Endpoint Data type Statistical methods Adult body mass Continuous GEEs, regression for repeated measures Adult feed consumption Continuous GEEs, regression for repeated measures Length of gestation Continuous Linear regression/anova Females whelping Binary Fisher s exact test Number kits whelped per female Count Negative binomial regression Number kits whelped live per female Count Negative binomial regression Kit mass at birth, three and six weeks of age Continuous Linear GEE regression Stillbirth and kit mortality at three and six weeks of age Binary Beta-binomial regression Change P in body P mass of seven-month-old juveniles during growth trial Continuous Linear GEE regression P PCBs and P TEQs in adult livers Continuous Linear regression/anova P PCBs and P TEQs in six-week-old kit livers Continuous Linear GEE regression PCBs and TEQs in seven-month-old juvenile livers Continuous Linear GEE regression PCB ¼ polychlorinated biphenyl; TEQ ¼ toxic equivalent; ANOVA ¼ analysis of variance; GEE ¼ generalized estimating equation.

5 784 Environ. Toxicol. Chem. 32, 2013 S.J. Bursian et al. probabilities is the joint probability of surviving to 31 weeks of age, and the complement (1 minus the joint probability) is the probability of death before 31 weeks of age. This method accounted for the kits necropsied postweaning. The survival models were beta-binomial models previously described for modeling mortality at three and six weeks of age. The LC50 was determined by iteratively solving for SPCBs. All statistical analyses were conducted using R statistical software ( including the additional R packages MASS for negative binomial models, aod for betabinomial models, geepack for GEE models, and doby for linear functions of estimated GEE regression parameters. RESULTS Dietary contaminant concentrations The analyzed concentrations of P PCBs in feed were similar to the nominal doses, and PCDDs and PCDFs contributed less than 2% P of the TEQs (Table 2). The analyzed concentrations of PCBs ( standard deviation) were (0.0016), 0.72 (0.12), 1.5 (0.21), 2.8 (0.38), 4.5 (0.49), and 6.1 (0.51) mg/g feed. These values will be used subsequently rather than the targeted concentrations. The corresponding concentrations of TEQs WHO 2005 ( standard deviation) were 0.41 (0.022), 4.8 (0.15), 10 (0.074), 18 (0.13), 28 (0.12), and 38 (0.38) pg/g feed, respectively. The percentage of contribution by PCDDs, PCDFs, non-ortho PCBs, and mono-ortho PCBs to the TEQs WHO 2005 in the treatment diets averaged 1.5, 1.4, 75, and 22%, respectively. Thus, 97% of the TEQs WHO 2005 in the diet was from PCBs with 3,3 0,4,4 0,5-pentachlorobiphenyl (PCB 126 using International Union of Pure and Applied Chemistry nomenclature) contributing 74% of the TEQs WHO The predominant organochlorine pesticide present in the feed was total toxaphene at a maximum concentration of 0.24 mg/g feed, followed by P dichlorodiphenyltrichloroethane (DDT) at a maximum concentration of mg/g feed (data not shown). Mercury was present at a maximum concentration of mg/g feed. The maximum P PBDE concentration was mg/g feed, with BDE 47 accounting for 73% of the P PBDE concentration (Supplemental Data, Table S1). Hepatic P PCB and TEQ WHO 2005 concentrations Hepatic P PCB and TEQ WHO 2005 concentrations of PCBexposed adult female mink were significantly greater (p < and p < 0.004, respectively) than control concentrations and increased monotonically with feed P PCB concentrations (Table 3). Concentrations in livers of adult males (not shown) were not significantly different from those in females. The percentage of contribution by PCDDs, PCDFs, non-ortho PCBs, and mono-ortho PCBs to the TEQs WHO 2005 in the maternal livers averaged 0.80, 1.4, 82, and 16%, respectively. Thus, 98% of hepatic TEQs WHO 2005 was from PCBs, with 82% being contributed by PCB 126. Hepatic P PCB and TEQ WHO 2005 concentrations (based only on non-ortho and mono-ortho PCBs) in livers of six-weekold kits (nine animals) and 31-week-old juveniles (23 animals) were significantly greater (p < 0.001) in the 0.72 mg P PCBs/g feed group compared to controls (15 kits and 30 juveniles; Table 4). The percentage of contribution by non-ortho and mono-ortho PCBs to the TEQs WHO 2005 averaged 77 and 23% in the kit livers and 81 and 19% in the juvenile livers, with PCB 126 contributing all of the non-ortho TEQs WHO Adult feed consumption Feed consumption of adult female mink did not differ significantly ( p ¼ 0.084) across treatment groups during the seven-week period prior to breeding. Average daily feed consumption for adult female mink was 115 g feed/d. Reproductive performance and offspring survivability Although all 75 adult females on the trial were successfully mated during the three-week breeding period, the mean number of kits whelped alive per litter was less at higher dietary PCB concentrations compared to lower dietary PCB concentrations (p < 0.004). The percentage of females whelping at least one kit (dead or alive) decreased slightly with increasing dietary PCB concentration, although this decrease was not statistically significant (p ¼ 0.110). The percentage of females whelping at least one kit was 100% for females in the control and and 1.5 mg P PCBs/g feed groups and 90, 87, and 73% for females in the 2.8-, 4.5-, and 6.1 mg P PCBs/g feed groups, respectively. One female each in the 2.8- and 4.5 mg P PCBs/g feed groups and three females in the 6.1 mg P PCBs/g feed group did not P whelp. Of these, one female each in the 2.8- and 6.1 mg PCBs/g feed groups had uterine implantation sites, indicating conception, while the female in the 4.5 mg P PCBs/g feed group and two females in the 6.1 mg P PCBs/g feed group did not. Average gestation lengths, which ranged from 46.3 to 48.7 d, were similar for all treatment groups (p ¼ 0.74). Compared to controls, the mean number of kits whelped alive per litter was 44% less (CI 17 63%, p ¼ ) in the 4.5 mg and 49% less (CI 21 67%, p ¼ ) in the 6.1 mg P PCBs/g feed groups (Table 5). The estimated relative rate of live kits per litter for a given increase in micrograms P PCBs/g feed (D) was e 0.17D (p < 0.001). Increases of 1 and 2 mg P PCBs/g feed were associated with 15.9 and 29.3% reductions in mean live kits whelped per litter. Kit mortality increased over time, ultimately resulting in no kits surviving to the end of the trial with the exception of those in the control and 0.72 mg P PCBs/g feed groups. The odds of kit mortality at six weeks of age were 22 (CI , p < 0.001) times greater in the 2.8 mg, 9.4 (CI , p ¼ ) times greater in the 4.5 mg, and 40 (CI , p ¼ ) times greater in the 6.1 mg P PCBs/g feed groups than the control group (Table 6). The numbers of kits alive at weaning, taking into consideration animals that were necropsied at six weeks of age, were 51, 30, 38, 9, 15, and 2 in the control and 0.72-, 1.5-, 2.8-, 4.5-, and 6.1 mg P PCBs/g feed groups, respectively. Between six and 10 weeks of age (when young mink were housed individually), 4 (7.8%), 6 (20%), 33 (87%), 8 (89%), 15 (100%), and 2 (100%) animals in the control and 0.72-, 1.5-, 2.8-, 4.5-, and 6.1 mg P PCBs/g feed groups died, respectively. The seven juvenile deaths that occurred from July 12, 2007, to the termination of the trial were one juvenile in the 0.72 mg P PCBs/g feed group, the remaining five juveniles in the 1.5 mg P PCBs/g feed group, and the remaining juvenile in the 2.8 mg P PCBs/g feed group. Thus, at the end of the trial, the only animals remaining were 47 juveniles in the control group and 23 juveniles in the 0.72 mg P PCBs/g feed group. Figure 2 depicts offspring mortality between 6 and 31 weeks of age, and Supplemental Data, Table S3 presents the number of offspring alive at key time points throughout the study. Body mass At the start of the experiment, adult female body mass did not differ significantly across treatment groups ( p ¼ 0.916) but

6 Effects of Hudson River fish on mink reproduction Environ. Toxicol. Chem. 32, Table 2. Mean concentrations of total polychlorinated biphenyls (PCBs), polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs), non-ortho PCBs, and mono-ortho PCBs as well as toxic equivalents (TEQs) a in experimental mink diets Dietary concentration (mg P PCBs/g feed) Compound TEF b Control PCDDs c 2,3,7,8-TCDD (ng/g) TEQs (pg/g) ,2,3,7,8-PeCDD (ng/g) TEQs (pg/g) ,2,3,4,7,8-HxCDD (ng/g) TEQs (pg/g) ,2,3,6,7,8-HxCDD (ng/g) TEQs (pg/g) ,2,3,7,8,9-HxCDD (ng/g) TEQs (pg/g) ,2,3,4,6,7,8-HpCDD (ng/g) TEQs (pg/g) OCDD (ng/g) PTEQs (pg/g) PCDDs (ng/g) PCDD TEQs (pg/g) %of P TEQs PCDFs c 2,3,7,8-TCDF (ng/g) TEQs (pg/g) ,2,3,7,8-PeCDF (ng/g) TEQs (pg/g) ,3,4,7,8-PeCDF (ng/g) TEQs (pg/g) ,2,3,4,7,8-HxCDF (ng/g) TEQs (pg/g) ,2,3,6,7,8-HxCDF (ng/g) TEQs (pg/g) ,2,3,7,8,9-HxCDF (ng/g) TEQs (pg/g) ,3,4,6,7,8-HxCDF (ng/g) TEQs (pg/g) ,2,3,4,6,7,8-HpCDF (ng/g) TEQs (pg/g) ,2,3,4,7,8,9-HpCDF (ng/g) TEQs (pg/g) OCDF (ng/g) PTEQs (pg/g) PCDFS (ng/g) PCDF TEQs (pg/g) %of P TEQs Non-ortho PCBs c PCB 77 (ng/g) d TEQs (pg/g) PCB 81 (ng/g) d TEQs (pg/g) PCB 126 (ng/g) d d d TEQs (pg/g) PCB 169 (ng/g) d d d d d d PTEQs (pg/g) Non-ortho PCBs (ng/g) Non-ortho PCB TEQs (pg/g) %of P TEQs Mono-ortho PCBs c PCB 105 (ng/g) d TEQs (pg/g) PCB 114 (ng/g) TEQs (pg/g) PCB 118 (ng/g) TEQs (pg/g) PCB 123 (ng/g) TEQs (pg/g) PCB 156 (ng/g) TEQs (pg/g) PCB 157 (ng/g) TEQs (pg/g) PCB 167 (ng/g) TEQs (pg/g) PCB 189 (ng/g)

7 786 Environ. Toxicol. Chem. 32, 2013 S.J. Bursian et al. Table 2. (Continued ) Dietary concentration (mg P PCBs/g feed) Compound TEF b Control PTEQs (pg/g) Mono-ortho PCBs (ng/g) Mono-ortho PCB TEQs (pg/g) %of P P TEQs PCBs (mg/g) c SD P TEQs (pg/g) SD % Lipids a TEQs based on nondetects ¼ 0.5 detection limit. b Toxic equivalency factors (TEFs) from Van den Berg et al. [31]. c Concentrations of PCDDs and PCDFs are means of three samples. Concentrations of PCBs and P PCBs are means of five samples. d Some values used in the calculation of these means were estimated as one-half of the detection limit. TCDD ¼ tetrachlorodibenzo-p-dioxin; PeCDF ¼ pentochlorodibenzo-p-dioxin; HxCDD ¼ hexachlorodibenzo-p-dioxin; HpCDD ¼ heptachlorodibenzo-pdioxin; OCDD ¼ octachlorodibenzo-p-dioxin; TCDF ¼ tetrachlorodibenzofuran; PeCDF ¼ pentachlorodibenzofuran; HxCDF ¼ hexachlorodibenzofuran; HpCDF ¼ heptachlorodibenzofuran; OCDF ¼ octachlorodibenzofuran. Nomenclature for the PCB congeners follows that of the International Union of Pure and Applied Chemistry: PCB 77 ¼ 3,3 0,4,4 0 -tetrachlorobiphenyl; PCB 81 ¼ 3,4,4 0,5-tetrachlorobiphenyl; PCB 126 ¼ 3,3 0,4,4 0,5-pentachlorobiphenyl; PCB 169 ¼ 3,3 0,4,4 0,5,5 0 -hexachlorobiphenyl; PCB 105 ¼ 2,3,3 0,4,4 0 -pentachlorobiphenyl; PCB 114 ¼ 2,3,4,4 0,5-pentachlorobiphenyl; PCB 118 ¼ 2,3 0,4,4 0,5- pentachlorobiphenyl; PCB 123 ¼ 2,3 0,4,4 0,5 0 -pentachlorobiphenyl; PCB 156 ¼ 2,3,3 0,4,4 0,5-hexachlorobiphenyl; PCB 157 ¼ 2,3,3 0,4,4 0,5 0 -hexachlorobiphenyl; PCB 167 ¼ 2,3 0,4,4 0,5,5 0 -hexachlorobiphenyl; PCB 189 ¼ 2,3,3 0,4,4 0,5,5 0 -heptachlorobiphenyl. generally decreased in all treatment groups over the first eight weeks of the trial. Body mass was significantly greater in the 1.5 mg (1,166 g, CI 1,119 1,213, p ¼ 0.014), 4.5 mg (1,190 g, CI 1,156 1,224, P p < 0.001), and 6.1 mg (1,206 g, CI 1,179 1,232, p < 0.001) PCBs/g feed groups compared to controls (1,103 g, CI 1,072 1,134). Body mass of individual kits was not significantly different at whelping (p ¼ 0.21), but average kit masses in the 1.5-, 2.8-, and 4.5 mg P PCBs/g feed groups (10, 19, and 28 TEQs WHO 2005/g feed, respectively) were, respectively, 27 g (CI , p ¼ ) or 25%, 29 g (CI , p ¼ 0.025) or 27%, and 50 g (CI , p < 0.001) or 46% less than controls at three weeks of age and 47 g (CI , p ¼ 0.017) or 21%, 47 g (CI , p ¼ 0.041) or 21%, and 73 g (CI , p ¼ ) or 33% less than controls at six weeks of age. The 6.1 mg P PCBs/g feed group was omitted from analysis at three and six weeks of age because the regression models were not estimable with only two live kits remaining in this treatment group (Supplemental Data, Table S4). The only significant treatment effect on juvenile growth was that males in the 0.72 mg P PCBs/g feed group gained 253 g (CI , p ¼ 0.003) or 34% more than males in the control group. At the end of the trial, mean change in body mass in the control group was 476 g for females and 750 g for males, and in the 0.72 mg P PCBs/g feed group mean change in body mass was 488 g for females and 1,003 g for males (data not presented). Lethal concentration estimates Estimated dietary and maternal hepatic LC20 and LC50 values based on kit stillbirths and kit mortality at three and six weeks of age were derived in terms of P PCBs/g and TEQs WHO 2005 /g (Supplemental Data, Table S5). Control adjusted LC20 and LC50 values were also estimated (Supplemental Data, Table S2). Plots illustrating dose response curves used to derive LC20 and LC50 values for kit mortality at six weeks of age based on dietary P PCB and TEQ WHO 2005 concentrations and maternal hepatic P PCB and TEQ WHO 2005 concentrations are presented in Figures 3 and 4, respectively. The estimated LC50 for mortality by 31 weeks of age was 0.78 mg P PCBs/g feed. DISCUSSION Dietary contaminant concentrations The analyzed concentrations of P PCBs ( P PCBs/g feed) in the treatment diets and the derived TEQs WHO 2005 ( pg TEQs WHO 2005 /g feed) were similar to those used in other mink feeding studies with a similar design [12 17]. In addition, the congener profile in most cases was similar across studies. Non-ortho PCBs (75%), and PCB 126 specifically (74%), contributed the majority of dietary TEQs WHO 2005 derived from HR fish. A mink feeding study of a similar design was conducted using PCB-contaminated fish collected from the Housatonic River (Massachusetts, USA) [16,17]. Housatonic River diets contained P PCB concentrations ranging from 0.34 to 3.7 mg/g feed and TEQ WHO 2005 concentrations ranging from 2.5 to 51 pg/g feed (recalculated using TEF values presented in Van den Berg et al. [31]), with 87% of the TEQs WHO 2005 being contributed by non-ortho PCBs and 81% by PCB 126 specifically. Two additional mink feeding studies of a comparable design used fish containing PCBs collected from Saginaw Bay (Lake Huron, Michigan, USA) or the mouth of the Saginaw River (Michigan, USA), which empties into Saginaw Bay. In the Saginaw Bay study [12 14], dietary concentrations ranged from 0.72 to 2.6 mg P PCBs/g feed (17 66 pg TEQs WHO 2005 /g feed, recalculated using TEF values presented in Van den Berg et al. [31]), with 64% of the TEQs WHO 2005 contributed by nonortho PCBs (62% by PCB 126). In the Saginaw River study [15], which was conducted 14 years after the Saginaw Bay study, dietary P PCB concentrations ranged from 0.83 to 1.7 mg P PCBs/g feed and 22 to 57 pg TEQsWHO 2005 /g feed (recalculated using TEF values presented in Van den Berg et al. [31]). In this study, non-ortho PCBs and PCDDs each contributed 35% of the TEQs WHO 2005 and PCDFs contributed 25%. The single largest contributor of TEQs WHO 2005 was PCB 126 (33%). It should be pointed out that there are some limitations with the TEQ approach. Toxic equivalency factors are consensus values of the relative potencies of the various PCB/PCDD/PCDF congeners rather than precise values. They may vary depending on species, measurement end points, and the relative proportions of individual congeners in complex mixtures. Some TEFs are based on in vitro studies that do not account for the potential

8 Effects of Hudson River fish on mink reproduction Environ. Toxicol. Chem. 32, Table 3. Mean toxic equivalents (TEQs) contributed by polychlorinated dibenzo-p-dioxins (PCDDs), polychlorinated dibenzofurans (PCDFs), non-ortho polychlorinated biphenyls (PCBs), and mono-ortho PCBs as well as P PCBs in livers of adult female mink fed diets containing contaminated fish from the Hudson River Feed group based on dietary concentration (mg P PCBs/g feed) Compound TEF a Control b TEQs contributed by PCDDs (pg/g wet wt) 2,3,7,8-TCDD c 0.21 c 0.24 c 0.24 c 0.36 c 0.52 c 1,2,3,7,8-PeCDD c 0.21 c 0.19 c 0.22 c 0.22 c 0.29 c 1,2,3,4,7,8-HxCDD c c c c c c 1,2,3,6,7,8-HxCDD c 0.18 c 0.23 c ,2,3,7,8,9-HxCDD c c c c c c 1,2,3,4,6,7,8-HpCDD OCDD P PCDD TEQs TEQs contributed by PCDFs (pg/g wet wt) 2,3,7,8-TCDF c c c c c c 1,2,3,7,8-PeCDF c c c c c c 2,3,4,7,8-PeCDF ,2,3,4,7,8-HxCDF c c c c ,2,3,6,7,8-HxCDF c c c ,2,3,7,8,9-HxCDF c c c c c c 2,3,4,6,7,8-HxCDF c c c c 1,2,3,4,6,7,8-HpCDF c c c c ,2,3,4,7,8,9-HpCDF c c c c c c OCDF P c c c c c c PCDF TEQs TEQs contributed by non-ortho PCBs (pg/g wet wt) c c P c c Non-ortho PCB TEQs TEQs contributed by mono-ortho PCBs (pg/g wet wt) P P Mono-ortho PCB TEQs TEQs (pg/g wet wt) d 2.4 (0.97) 33 (16) 61 (25) 101 (33) 181 (55) 220 (63) P PCBs (mg/g wet wt) d (0.032) 1.5 (0.86) 2.9 (1.8) 3.4 (2.2) 5.0 (3.7) 6.2 (4.8) % Lipid (95% CI) 12.5 ( ) 11.0 ( ) 12.2 ( ) 9.1 ( ) 8.5 ( ) 9.5 ( ) a Toxic equivalency factors (TEFs) from Van den Berg et al. [31]. b Values are TEQs based on nondetects ¼ 0.5 detection limit. c Some values used in the calculation of these means were estimated as one-half of the detection limit. d Number in parentheses is the standard deviation. TCDD ¼ tetrachlorodibenzo-p-dioxin; PeCDF ¼ pentochlorodibenzo-p-dioxin; HxCDD ¼ hexachlorodibenzo-p-dioxin; HpCDD ¼ heptachlorodibenzo-pdioxin; OCDD ¼ octachlorodibenzo-p-dioxin; TCDF ¼ tetrachlorodibenzofuran; PeCDF ¼ pentachlorodibenzofuran; HxCDF ¼ hexachlorodibenzofuran; HpCDF ¼ heptachlorodibenzofuran; OCDF ¼ octachlorodibenzofuran. Nomenclature for the PCB congeners follows that of the International Union of Pure and Applied Chemistry: PCB 77 ¼ 3,3 0,4,4 0 -tetrachlorobiphenyl; PCB 81 ¼ 3,4,4 0,5-tetrachlorobiphenyl; PCB 126 ¼ 3,3 0,4,4 0,5-pentachlorobiphenyl; PCB 169 ¼ 3,3 0,4,4 0,5,5 0 -hexachlorobiphenyl; PCB 105 ¼ 2,3,3 0,4,4 0 -pentachlorobiphenyl; PCB 114 ¼ 2,3,4,4 0,5-pentachlorobiphenyl; PCB 118 ¼ 2,3 0,4,4 0,5- pentachlorobiphenyl; PCB 123 ¼ 2,3 0,4,4 0,5 0 -pentachlorobiphenyl; PCB 156 ¼ 2,3,3 0,4,4 0,5-hexachlorobiphenyl; PCB 157 ¼ 2,3,3 0,4,4 0,5 0 -hexachlorobiphenyl; PCB 167 ¼ 2,3 0,4,4 0,5,5 0 -hexachlorobiphenyl; PCB 189 ¼ 2,3,3 0,4,4 0,5,5 0 -heptachlorobiphenyl. differences between animal species in absorption, distribution, metabolism, and elimination. Other TEFs are based on quantitative structure activity relationships because of a lack of toxicity data for some congeners, which can introduce a degree of uncertainty into the determination of TEQs. As such, TEFs may under- or overestimate the relative potencies of congeners and, thus, should be considered as protective estimates as opposed to predictors of effect thresholds [20]. The dietary concentrations of P PCBs used in the present study bracket dietary concentrations of P PCBs that are relevant for wild mink residing along the upper Hudson River. The National Oceanic and Atmospheric Administration (NOAA) database of tissue residue data for specific watershed projects was searched for whole-body analyses of fish sampled in the Hudson River between latitudes and , which encompasses the three carp collection sites for the present study and fish sizes of 7 to 20 cm, which is the expected size of typical prey fish consumed by mink [17]. The data used were from sampling studies conducted by NOAA in 1993, 1995, and 1999 and by the General Electric Company from 2004 to Fish species sampled in this region included the largemouth bass (Micropterus salmoides), pumpkinseed sunfish (Lepomis gibbosus), redbreast sunfish (Lepomis auritus), yellow perch (Perca flavescens), golden shiner (Notemigonus crysoleucas), bluntnose minnow (Pimephales notatus), spottail shiner (Notropis hudsonius), common shiner (Luxilus cornutus), spotfin

9 788 Environ. Toxicol. Chem. 32, 2013 S.J. Bursian et al. Table 4. Mean concentrations of total polychlorinated biphenyls (PCBs), non-ortho PCBs, and mono-ortho PCBs as well as toxic equivalents (TEQs) in livers of six-week-old kits and 31-week-old juvenile mink fed diets containing contaminated fish from the Hudson River Feed group based on dietary concentration (mg P PCBs/g feed) Control a 0.72 Control a week-old kits 31-week-old juveniles Compound TEF b (ng/g wet wt) Concentration TEQ Concentration (pg/g wet wt) (ng/g wet wt) TEQ Concentration (pg/g wet wt) (ng/g wet wt) TEQ Concentration (pg/g wet wt) (ng/g wet wt) TEQ (pg/g wet wt) Non-ortho PCBs c c c c P c d c P Non-ortho PCBs, TEQs Mono-ortho PCBs P P Mono-ortho PCBs, TEQs PCBs (mg/g wet wt) P TEQs (pg/g wet wt) e (0.017) 1.7 (0.85) 1.2 (0.99) 39 (39) (0.021) 1.2 (0.63) 0.89 (0.28) 23 (5.7) % Lipid (95% CI) Males: 11.5 ( ) Females: 12.3 ( ) Males: 7.8 ( ) Females: 11.4 ( ) Males: 14.4 ( ) Females : 14.5 ( ) Males: 14.0 ( ) Females: 14.2 ( ) a Values are based on nondetects ¼ 0.5 detection limit. b Toxic equivalency factors (TEFs) from Van den Berg et al. [31]. c Some values used in the calculation of these means were estimated as one-half of the detection limit. TCDD ¼ tetrachlorodibenzo-p-dioxin; PeCDF ¼ pentochlorodibenzo-p-dioxin; HxCDD ¼ hexachlorodibenzo-p-dioxin; HpCDD ¼ heptachlorodibenzo-pdioxin; OCDD ¼ octachlorodibenzo-p-dioxin; TCDF ¼ tetrachlorodibenzofuran; PeCDF ¼ pentachlorodibenzofuran; HxCDF ¼ hexachlorodibenzofuran; HpCDF ¼ heptachlorodibenzofuran; OCDF ¼ octachlorodibenzofuran. Nomenclature for the PCB congeners follows that of the International Union of Pure and Applied Chemistry: PCB 77 ¼ 3,3 0,4,4 0 -tetrachlorobiphenyl; PCB 81 ¼ 3,4,4 0,5-tetrachlorobiphenyl; PCB 126 ¼ 3,3 0,4,4 0,5-pentachlorobiphenyl; PCB 169 ¼ 3,3 0,4,4 0,5,5 0 -hexachlorobiphenyl; PCB 105 ¼ 2,3,3 0,4,4 0 -pentachlorobiphenyl; PCB 114 ¼ 2,3,4,4 0,5-pentachlorobiphenyl; PCB 118 ¼ 2,3 0,4,4 0,5- pentachlorobiphenyl; PCB 123 ¼ 2,3 0,4,4 0,5 0 -pentachlorobiphenyl; PCB 156 ¼ 2,3,3 0,4,4 0,5-hexachlorobiphenyl; PCB 157 ¼ 2,3,3 0,4,4 0,5 0 -hexachlorobiphenyl; PCB 167 ¼ 2,3 0,4,4 0,5,5 0 -hexachlorobiphenyl; PCB 189 ¼ 2,3,3 0,4,4 0,5,5 0 -heptachlorobiphenyl. shiner (Notropis spilopterus), and mimic shiner (Notropis volucellus). The average P PCB concentration for all fish sampled between these two latitudes was 4.0 mg/g tissue (n ¼ 430, range ). Fish collected at sites closest to the carp collection sites in the present study had average P PCB concentrations of 4.6 mg/g tissue for the Moses Kill area (n ¼ 55, range ), 4.5 mg/g tissue for the Northumberland Pool area (n ¼ 36, range ), and 2.4 mg/g tissue for the above Lock 2 site (n ¼ 70, range ). The average P PCB concentration for fish sampled at these three sites on the Hudson River was 3.8 mg/g tissue (Supplemental Data, Table S6). Daily doses (based on dietary concentrations and an average daily feed consumption of 97 g/kg body mass that was calculated by dividing average feed consumption per animal [115 g] by average body wt of adult females across the six treatment groups [1,186 g]) of the various organochlorine pesticides as Table 5. Mean number of kits whelped alive per litter from mink fed diets containing polychlorinated biphenyl (PCB) contaminated fish collected from the Hudson River a 95% CI 95% CI Feed group based on dietary concentration Mean live kits whelped per litter (68) b LCL UCL Difference (%) in rate from control LCL UCL p c Control NA NA NA NA 0.72 mg P PCBs/g feed mg P PCBs/g feed mg P PCBs/g feed mg P PCBs/g feed mg P PCBs/g feed a Treatment effects were estimated with negative binomial regression models. b Results based on 68 litters with at least one kit born dead or alive. c p value for comparison to control group. CI ¼ confidence interval; LCL ¼ lower confidence limit; UCL ¼ upper confidence limit.

10 Effects of Hudson River fish on mink reproduction Environ. Toxicol. Chem. 32, Table 6. Mortality of kits whelped and nursed by mink fed diets containing polychlorinated biphenyl (PCB) contaminated fish collected from the Hudson River a Litter-adjusted 95% CI End point Feed group based on dietary concentration Kit mortality (%) b Kit mortality (%) LCL UCL p c Kit mortality at three weeks ( p < 0.001) d Control NA 0.72 mg P PCBs/g feed mg P PCBs/g feed mg P PCBs/g feed < mg P PCBs/g feed mg P PCBs/g feed <0.001 Kit mortality at six weeks (p < 0.001) d Control NA 0.72 mg P PCBs/g feed mg P PCBs/g feed mg P PCBs/g feed < mg P PCBs/g feed mg P PCBs/g feed a Treatment effects were estimated with beta-binomial regression models. b Results based on 321 kits born live from 61 litters with a least one live birth. c p value for comparison to control group. d p value for overall test of treatment effect. CI ¼ confidence interval; LCL ¼ lower confidence limit; UCL ¼ upper confidence limit; NA ¼ not available. well as the potentially toxic metals (arsenic, cadmium, chromium, copper, mercury, nickel, lead, selenium, and zinc) present in the treatment diets were less than the toxicity reference values for mink reported by Hink et al. [35]. Similarly, dietary concentrations of PBDEs were less than those reported to cause adverse effects in mink [36]. Hepatic P PCB and TEQ WHO 2005 concentrations Hepatic concentrations of P PCBs and TEQ WHO 2005 concentrations increased with dietary concentration and were similar between adult male and female mink. The percentage of hepatic TEQs WHO 2005 compared to dietary TEQs WHO 2005 contributed by non-ortho PCBs increased slightly (from 75 to 82%), while the contribution by mono-ortho PCBs decreased by 5%. The PCB 126 contribution increased from 74 to 82% when comparing dietary to hepatic TEQs WHO 2005, respectively. Bioaccumulation factors (BAFs) were derived by dividing the hepatic concentration (wet wt) of P PCBs and TEQs WHO 2005 by dietary P PCB and TEQ concentrations (wet wt), respectively. The BAF for P PCBs was 1.2 and the BAF for TEQs WHO 2005 was 5.9. These BAFs are similar to those based on dietary and hepatic P PCB and TEQ WHO 2005 concentrations reported Fig. 2. Number of live kits in each feed group based on dietary concentration at 6, 10, and 31 weeks of age. The number of live kits at six weeks of age reflects the number alive after necropsy. The percentage of live kits at 10 and 31 weeks of age compared to six weeks of age is given in parentheses.

11 790 Environ. Toxicol. Chem. 32, 2013 S.J. Bursian et al. Fig. 3. Lethal concentration for 50% (LC50) and 20% (LC20) with 95% confidence interval (horizontal interval) for increased mortality of six-week-old kits based on micrograms P polychlorinated biphenyls (PCBs)/g feed (top) and picograms toxic equivalents (TEQs)/g feed (bottom). for the Housatonic River study (1.2 and 4.0, respectively) [16,17]. The BAFs for individual PCB congeners that contributed P the majority of TEQs WHO 2005 resembled the BAF for TEQsWHO 2005 (PCB 126 ¼ 6.5, PCB 105 [2,3,3 0,4,4 0 -pentachlorobiphenyl] ¼ 4.1, PCB 118 [2,3 0 4,4 0,5- pentachlorobiphenyl] ¼ 4.3, and PCB 156 [2,3,3 0,4,4 0,5- hexachlorobiphenyl] ¼ 7.7). Hepatic concentrations of P PCBs and TEQs WHO 2005 in six-week-old kits and 31-week-old juveniles were similar to adult concentrations, as was reported for mink exposed to PCBs derived from fish collected from the Housatonic River [17]. Adult feed consumption Adult feed consumption did not differ significantly over the first seven weeks of the trial, although a flavor enhancer was added to the feed beginning in week 6 following the observation during week 5 that feed consumption decreased by approximately 50% in those groups fed diets containing primarily herring (control and and 1.5 mg P PCBs/g feed groups). The inclusion of a flavor enhancer in all of the treatment diets resulted in an immediate increase in feed intake such that feed consumption was again comparable across all groups by week 6. The average daily feed consumption of 115 g reported here is identical to the value reported for adult female ranch mink [37] and similar to the 127 g/d rate reported for mink fed diets containing fish collected from the Housatonic River [16]. Reproductive performance and offspring mortality The effects of consumption of diets containing PCBcontaminated fish collected from the Hudson River on reproductive performance were similar to those reported for mink fed diets containing PCB/PCDD/PCDF-contaminated fish collected from Saginaw Bay [12,14]. There was no significant effect on the percentage of females whelping at least one kit, but there was significant reduction in the mean number of kits whelped alive per litter in both studies. The no observed adverse effect levels (diet-based)/concentrations (tissue-based; NOAELs/ NOAECs) and lowest observed adverse effect levels/concentrations (LOAELs/LOAECs) are presented in Table 7. An increase in kit mortality is the effect most often reported in mink reproduction studies using PCB/PCDD/PCDFcontaminated fish. In addition to the present study, kit mortality was significantly increased in the Saginaw Bay study [12,14] and the Housatonic River study [16]. In the latter study, the contribution by non-ortho PCBs and PCB 126 to P TEQs was similar to the present study. The NOAEL/NOAEC and LOAEL/ LOAEC for each study are presented in Table 7. It is interesting to note that an equivalent dietary concentration of TEQs WHO 2005 provided only by PCB 126 had no effect on kit mortality through six weeks of age [18]. The differences in NOAEL(C)s ranged from 3.6-fold (hepatic TEQs) to 106-fold (dietary P PCBs). The differences in LOAEL(C)s for increased kit mortality were considerably less, ranging from 1.5-fold (hepatic P PCBs) to fivefold (dietary P PCBs).

12 Effects of Hudson River fish on mink reproduction Environ. Toxicol. Chem. 32, Fig. 4. Lethal concentration for 50% (LC50) and 20% (LC20) with 95% confidence interval (horizontal interval) for increased mortality of six-week-old kits based on maternal hepatic concentrations of P polychlorinated biphenyls (PCBs; micrograms per gram liver, wet wt; top) and toxic equivalents (TEQs; picograms per gram liver, wet wt; bottom). A treatment-related increase in juvenile mortality between weaning at six weeks of age and the end of the growth period at approximately 30 weeks of age was reported only in the present study. Between 6 and 10 weeks of age, siblings were still grouphoused and the larger, stronger animals were more effective at competing for feed compared to their smaller, weaker littermates. Smaller animals that continue to grow poorly may die, even after being housed singly at 10 to 12 weeks of age. Because dietary concentrations of 1.5 mg P PCBs/g feed and greater resulted in decreased kit body mass, it was not unexpected that animals in these treatment groups experienced excessive mortality after weaning. In the Saginaw Bay study [12,14], kits were not maintained beyond weaning. In the Housatonic River study [16], offspring maintained through 27 weeks had similar growth rates. Body mass The significantly greater body masses of adult females at the higher dietary concentrations of PCBs compared to controls could be a reflection of the initial decrease in feed intake of control animals fed diets containing 20% ocean herring. While feed consumption was equivalent across groups after beef flavoring was added to all treatment diets, it is apparent that control females continued to have a lower body mass. Body masses of treated kits were comparable at whelping but significantly less compared to controls at six weeks of age in both the Saginaw Bay study [12,14] and the present study. The NOAELs/NOAECs and LOAELs/LOAECs for the respective studies are presented in Table 7. In contrast, body mass of kits exposed to an equivalent or greater concentration of TEQs WHO 2005 provided only by PCB 126 was not adversely affected [18]. It is not clear why male juveniles in the 0.72 mg P PCBs/g feed group gained 253 g more mass compared to controls by the end of the trial. In the Housatonic River study [16], exposure to PCBs had no effect on juvenile growth through 27 weeks of age. Lethal concentration estimates Because there are acknowledged limitations associated with use of NOAELs/NOAECs and LOAELs/LOAECs [37,38] in risk assessment, LC20 and LC50 values were derived for kit mortality in the present study. Regression estimates of lethal concentrations (LC20 and LC50) are less sensitive to the idiosyncrasies of experimental design, and estimates are not limited to the concentrations under study. The only mink feeding study of a similar design as the present study that reported lethal dietary concentrations was the Housatonic River study [16]. There is a 2.9-fold difference between the Hudson River and Housatonic River LC20 values based on P PCBs with overlapping confidence intervals. The LC20 values based on TEQs WHO 2005 differ by 5.5-fold with no overlap in confidence intervals (Table 7). Assuming an average daily feed intake of 97 g/kg body mass (based on average feed consumption and body weight of adult females across the six treatment groups), the dose lethal to 20% of the population (LD20) was